The control of nitrous oxide emissions from agricultural andnatural soils

U. Skiba a,*, K.A. Smith b

a Institute of Terrestrial Ecology, Bush Estate, Penicuik, Midlothian, EH26 0QB Scotland, UKb Institute of Ecology and Resource Management, University of Edinburgh, West Mains Road, Edinburgh, EH9 3JG Scotland, UK

Received 20 July 1999; accepted 22 November 1999

Importance of this paper: This paper reviews the available information on how emissions of N2O from soils are con-

trolled by the interaction of parameters such as substrate (N) supply, soil water content and temperature. It shows how

these interactions help to explain observed deviations in emissions from values predicted by current IPCC methodology,

which is based solely on N inputs.

Abstract

This paper provides a summary of our current understanding of the key drivers of N2O emissions from soil in

temperate and tropical, natural and agricultural ecosystems. These drivers are substrate supply, as N additions and

mineralisation of organic N in soil, soil water content and temperature. They can exert synergistic or antagonistic

in¯uences on the emissions which can vary spatially and temporally. Such in¯uences explain why emission rates often

di�er greatly from those based on current IPCC methodology. The latter only takes account of N inputs: direct

emissions from agricultural soils are taken to be 1.25% of the N applied, while those from natural soils are taken to be

1% of the N deposited from the atmosphere, however, observed values range from 0.2% to 15%. Inadequate accounting

for all sources a�ecting levels of soil mineral N (e.g. freeze±thaw cycles, ploughing, biomass burning, the ®rst rainfall in

wet seasons) and inter-annual di�erences in the size and timing of rainfall events in relation to land management

practices are prime causes of the deviations. Ó 2000 Elsevier Science Ltd. All rights reserved.

Keywords: Agricultural soils; Natural soils; N fertiliser; N deposition; Soil water; Temperature

1. Introduction

The soil is the dominant source of atmospheric ni-

trous oxide (N2O) contributing about 57% (9 Tg yÿ1) of

the total annual global emission (IPCC, 1997). Increas-

ing N inputs to agricultural and natural soils have

greatly increased emissions from this source in the last

few decades (Kroeze et al., 1999). Current IPCC meth-

odology for calculating national N2O emission inven-

tories relates the emissions to N input by mineral

fertilisers, animals and N deposition. It is assumed that

1.25% of the N applied as synthetic fertiliser, manure or

crop residues to agricultural land and 1% of the N

supplied by atmospheric deposition to natural soils is

emitted as N2O (IPCC, 1997). These simple estimates

(Ôdefault valuesÕ) have their merits, as they are based on

readily available input data, but they are unable to

capture the complex interaction of the soil processes and

variables that control N2O emission. The key variables

leading to N2O emission are well understood and have

been reviewed extensively by e.g. Bouwman (1990),

Williams et al. (1992) and Granli and Bùckman (1994),

and some aspects are discussed in detail elsewhere in this

Chemosphere ± Global Change Science 2 (2000) 379±386

* Corresponding author. Tel.: +1-44-131-445-8532; fax: +1-

44-131-445-3943.

E-mail address: ums@ite.ac.uk (U. Skiba).

1465-9972/00/$ - see front matter Ó 2000 Elsevier Science Ltd. All rights reserved.

PII: S 1 4 6 5 - 9 9 7 2 ( 0 0 ) 0 0 0 1 6 - 7

issue. The general consensus of these reviews is that the

rate of N2O production and emission primarily depends

on the availability of a mineral N source (substrate for

nitri®cation or denitri®cation), and on soil temperature,

soil water content, and (for denitri®cation) the avail-

ability of labile organic compounds. These variables are

universal, but operate in di�erent combination and or-

der of importance in both space and time. We outline

here our current knowledge on how these variables in-

teract to control N2O emissions from agricultural and

natural soils in temperate and tropical climates, in the

hope that they will facilitate the design of future exper-

iments to quantify ¯ux-variable relationships, and the

development of predictive models.

2. Processes

Two biological processes with contrasting require-

ments for oxygen are the principal sources of N2O and

contribute some two-thirds of all emissions (Bremner

and Blackmer, 1981; Payne, 1981): Nitri®cation is an

aerobic process in which ammonium (NH�4 ) is oxidised

to nitrate (NOÿ3 ) (Robertson, 1989). At sub-optimal

oxygen concentrations oxidation to NOÿ3 is incomplete

and some of the NH�4 is channelled into production of

NO and N2O (Poth and Focht, 1985). Most commonly,

nitri®cation is carried out by very speci®c autotrophic

organisms, which are able to grow at the expense of

generating energy from nitri®cation. The best studied

are the obligate chemoautrotrophs, Nitrosomonas and

Nitrobacter species (Robertson and Kuenen, 1991). A

range of heterotrophic bacteria, fungi and algae can also

nitrify and produce N2O, however, usually at much

slower rates (Killham, 1986). The contribution from the

di�erent groups varies widely. For example, in a mixed

woodland on a silt loam, chemotrophic nitri®cation was

responsible for 50% of the total N2O emission (Peter-

john et al., 1998), while in an acid forest soil 8% of the

observed nitri®cation rate was attributed to the oxida-

tion of organic N to NOÿ3 (Barraclough and Puri, 1995);

and in a deciduous woodland soil fungal nitri®cation

appeared to be the main source of N2O (Castaldi and

Smith, 1998).

Denitri®cation is the anaerobic reduction of NOÿ3 to

N2O and N2 (Payne, 1981). A wide range of bacteria are

able to denitrify. They are facultative anaerobes and

switch to NOÿ3 as terminal electron acceptor when the

O2 concentrations in soil become depleted. The ratio of

N2O to N2 production depends on the species of denit-

ri®er involved (Robertson and Kuenen, 1991), on the

degree of anaerobicity in soil, soil carbon and NOÿ3content and soil pH, as discussed further by e.g.

Sahrawat and Keeney (1986) and Granli and Bùckman

(1994). The largest rates of N2O emission tend to be

associated with denitri®cation. Rates of N2O production

by nitri®cation tend to be smaller (Williams et al., 1992);

however, soil conditions favourable for nitri®cation to

proceed are much more common, so that the contribu-

tion of nitri®cation to the total global N2O emission

may not be trivial.

The balance between the two processes, nitri®cation

and denitri®cation, contributing to the N2O emission

will vary with climate, soil conditions and soil manage-

ment. Generally, high rainfall, poor drainage, ®ne soil

texture and high organic carbon content promote deni-

tri®cation and associated N2O production, whereas low

rainfall, good drainage and aeration and coarse texture

promote nitri®cation and associated N2O production

(Gro�man, 1991). However, in most soils the prevalence

of nitri®cation or denitri®cation as the main source of

N2O is not static and can switch very rapidly, as the soil

aeration state within the biologically active site changes

due to e.g. rainfall or increased O2 demand caused by

the presence of easily mineralisable organic matter.

3. Variables controlling the rate of N2O emission

3.1. Mineral N

The availability of mineral N as a substrate for ni-

tri®cation and denitri®cation is an essential requirement

for stimulating N2O emission. The e�ect of mineral N

additions on the N2O ¯ux has been studied and reviewed

extensively (Bouwman, 1990; Eichner, 1990; Granli and

Bùckman, 1994; Mosier, 1994), particularly in inten-

sively managed agricultural soils in developed countries

in the temperate climate zone. Provided there is a pres-

ence of an active nitrifying and denitrifying population

and soil aeration is optimal for N2O production by ni-

tri®cation or denitri®cation, application of mineral N

fertilisers increases N2O emissions very rapidly. Am-

monium-based fertilisers and urea (after hydrolysis to

NH3) provide substrates for nitri®cation, and the

product of nitri®cation, NO3, will provide a substrate

for denitri®cation. Thus N2O may be derived from NH4-

based fertilisers and urea during nitri®cation or through

subsequent denitri®cation, whereas NO3-based fertilisers

provide substrates for denitri®cation only. The increases

in nitri®cation and denitri®cation rates and accompa-

nying N2O losses tend to be short-lived, and near-

background emissions are restored when the substrate is

depleted, which usually occurs within a few days to

several weeks after application. This typical pattern has

been observed for fertilised grassland in the tropics

(Veldkamp et al., 1998) and in temperate western

Europe (Velthof et al., 1996; Dobbie et al., 1999) and

in arable crops (e.g. Jacinthe and Dick, 1997).

Equally, additions of excreta and urine of grazing

animals induce a similar response in N2O emissions. In

the soil the organic N in faeces and urine is rapidly

380 U. Skiba, K.A. Smith / Chemosphere ± Global Change Science 2 (2000) 379±386

mineralised to NH4. Nitrous oxide emissions from urine

patches peaked around 10±15 days after deposition

(Flessa et al., 1996; Yamulki and Jarvis, 1997). For dung

patches the timing was more variable: from 10 days after

deposition (Flessa et al., 1996) to 4 months after depo-

sition (Yamulki and Jarvis, 1997).

Current estimates of global N2O emissions calculate

the direct emissions of N2O from agricultural soils as a

®xed percentage of the mineral or organic N applied, 1.25

(0.25±2.25)% (IPCC, 1997). This default emission value

has been obtained from 20 studies of one year long ¯ux

measurements in several countries, mainly Germany,

Japan, UK and USA, (Bouwman, 1994). Since then

more recent studies have shown considerable departures

from the default. For example; lower emission values

were calculated for wheat and oil seed rape ®elds in

Central France (0.42±0.55%) (Henault et al., 1998) and a

wheat ®eld in South England (0.93%) (Yamulki et al.,

1995). For corn±soyabean±alfalfa rotations in Canada,

emission factors ranged from 1% to 4% (Mackenzie et

al., 1998), for arable crops in Germany from 0.7 to 8.5

(Kaiser et al., 1998) and for wheat ®eld in Mexico from

1.7% to 3.8% (Ortiz-Monasterio et al., 1996). Very high

emission factors were measured from a potato crop in

Germany, where 10.7±16% of the N applied to the po-

tato crop were emitted as N2O (Ruser et al., 1998). This

variation in fertiliser-induced emission factors and de-

viation from the 1.25% IPCC default value is further

exempli®ed by work by one of our groups in Scotland

(Fig. 1(a)). Emission factors from fertilised arable and

grassland crops ranged from 0.2% to 0.7% for wheat

and barley, 1.8±7% for potato and brassica crops and

0.3±5.8% for fertilised grassland (Dobbie et al., 1999).

The latter study not only shows large deviations from

IPCC default values, but also suggests that in some

situations the type of agricultural crop can signi®cantly

in¯uence the rate of N2O emissions per unit of N applied

and that inter-annual variations in N2O emissions can

be large. The inter-annual variations are exempli®ed in

Fig. 1(a) by the variations in grassland N2O emissions

which received 300±360 kg N haÿ1 yÿ1 over a six-year

period. The total annual N2O loss varied from 0.4% to

5.8% of the N input (Dobbie et al., 1999). In a study in

Germany, where N2O ¯uxes were measured over a three-

year period from cereal and arable crops, inter-annual

variations were also high, and as in the Scottish study,

emission losses were highest for the brassica crop and

sugar beet (1.5±4.1%), followed by oilseed rape (1.4±

2.5%) and winter barley (1.4±2.3%) (Kaiser et al., 1998).

The ranges in brackets are the minimum and maximum

values for individual years and demonstrate the inter-

annual variations.

The literature, however, is divided on the importance

of crop type in determining the rate of N2O emission.

For instance in western Europe N2O emissions were

generally larger from potatoes, sugar beet and broccoli

than from wheat, barley or oilseed rape (Henault et al.,

1998; Kaiser et al., 1998; Smith et al., 1998; Dobbie et al.,

1999) and data from Colorado, USA, have shown

comparable di�erences between small grains and maize

(Mosier, personal communication). However, a wider

comparison including data from many regions did not

reveal any such e�ect overall (Mosier and Kroeze, 1999).

Nitrous oxide emissions from grasslands can be higher

than for many arable crops and a mean emission factor

of 2% was estimated for a range of grasslands in Europe

(Fowler et al., 1997).

The reasons for di�erences in N2O emissions ac-

cording to crop type are primarily related to the crop

requirements for speci®c climatic conditions and man-

agement. For example, in temperate climates intensively

Fig. 1. Deviations of N induced emissions of N2O in Scotland

from the IPCC Ôdefault valuesÕ. (a) mineral N induced emissions

from grass (�), arable-cereal ( ), arable-non cereal (h), IPCC

emission factor 1.25% (solid line). Data from Dobbie et al.

(1999); (b) N deposition induced emissions from forest and

moorland soils, in upland areas ( ), large-scale acid mist ex-

periments (s), and downwind of poultry and pig farms (�),IPCC emission factor 1% (solid line). Data from Skiba et al.

(1998).

U. Skiba, K.A. Smith / Chemosphere ± Global Change Science 2 (2000) 379±386 381

managed grasslands tend to be concentrated in high

rainfall regions, often on soils too wet for arable crops.

Grazing animals compact the soil, and together the

wetness and compaction provide suitable conditions for

N2O production. The timing of fertiliser application, the

soil water removal by crops and/or the quantities of

plant residues all in¯uence soil N and water dynamics

which e�ect N2O production and emission rates. When

di�erent crop types were not associated with di�erent

soil conditions no crop e�ect was observed: In Costa

Rica N2O emissions from pasture, sugar cane and ba-

nana plantations were completely independent of crop

type; emissions were solely controlled by the soil water

content (Veldkamp et al., 1998). The same study showed

that replacing a traditional pasture, dominated by the

grass Ischaeum indicum, with a grass/legume mixture of

Brachiaria decumbens and Arachis pintoi improved dry

matter yield without increasing emissions of N2O. In

Belize, the soil drainage status outweighed any e�ects of

vegetation type (grass, arable, forest or cleared forest)

on the emissions of N2O (Rees et al., 1996).

The inter-annual variations in the German and

Scottish studies have di�erent origins. In the latter the

higher annual emissions occurred when rainfall during

the summer months was high. In the German study

di�erences in freeze±thaw cycles in winter and rainfall

events in summer contributed to the inter-annual vari-

ations. The e�ects of freeze±thaw cycles and rainfall are

very di�erent: rainfall will provide the necessary anaer-

obic microsites for N2O production, but freeze±thaw

cycles stimulate mineral N production and thereby in-

crease N2O. The general assumption used to be that

N2O emissions during cold winter periods are very low

and therefore year-round measurements were usually

much less intensive during this cold period. Recent

studies, however, have shown that freeze±thaw cycles

(R�over et al., 1998) and periods of low soil temperatures

(<4°C) can increase emissions signi®cantly, this is

thought to be due to the stimulation of microbial ac-

tivity as a result of additional available C and N from

microbial cells killed by freezing or low temperatures

(Christensen and Tiedje, 1990). This phenomenon is not,

strictly speaking, a temperature e�ect but a N e�ect, by

increasing N mineralisation. Such emissions from arable

land have been shown to account for 50% of the total

annual N2O ¯ux (Kaiser et al., 1998). N2O emissions

have even been measured through snow cover in agri-

cultural and also pristine alpine ecosystems (Van

Bochove et al., 1996).

In climates with distinct wet and dry seasons, the ®rst

rainfall onto dry soil can have a very similar e�ect to

that of freeze±thaw cycles and has been shown to create

a large pulse of CO2, NO and N2O emissions. These

pulses of trace gas emissions are thought to be caused by

accumulation of inorganic N in dry soils and reactiva-

tion of water-stressed bacteria upon wetting, which then

metabolize the pool of available inorganic N (Cabrera,

1993; Davidson et al., 1993). For a seasonally dry

tropical forest soil in Mexico, experimental wetting of

the ®eld plots at the end of the dry season resulted in

rapid and large pulses of NO, N2O and CO2 (Davidson

et al., 1993). Pulsing is not necessarily con®ned to

tropical countries; in a dry rye stubble ®eld in Denmark,

simulated rainfall also increased emissions of N2O

(Jùrgensen et al., 1998).

3.2. N deposition

The wet deposition of NOÿ3 and NH�4 and the dry

deposition of NO2 and NH3 can provide a signi®cant

source of N to the non-agricultural, non-fertilised soil

and therefore a substrate for nitri®cation and deni-

tri®cation and associated releases of gaseous forms of

N. Globally, atmospheric N deposition rates range

from <2 kg N haÿ1yÿ1 in pristine environments to

>80 kg N haÿ1 yÿ1 immediately adjacent to large point

sources of pollution, such as industrialised conurbations

(NOx) or intensive livestock enterprises (NH3) (Fowler

et al., 1998). N deposition rates may also be enhanced by

altitude and trees, due to the frequent occurrence of hill

clouds, which carry the major ions in a much more

concentrated form than rain water. Upland forests are

very e�cient in capturing cloud water, and the rates of

N deposition to forests may be a factor of two larger

than to other ecosystems (Fowler et al., 1989, 1998).

Thus in some upland areas of Northern Britain and

South Germany, for example, N deposition rates of

40 kg N haÿ1 yÿ1 have been measured (Aber et al., 1989;

Fowler et al., 1989; Rennenberg et al., 1998). In contrast

to fertiliser additions, N deposition provides a continu-

ous small elevated input of N. Experimental simulations

of such additions have shown, that initially the extra N

applied becomes immobilised for many years, but

eventually these additions will exceed demand and in-

creased rates of N leaching and gaseous N emission can

be expected (Aber et al., 1989; Bowden et al., 1991;

Skiba et al., 1999). This dual response is shown in

Fig. 1(b). Data points that lie signi®cantly above the 1%

IPCC emission factor came from locations close to point

sources (poultry and pig farms), which had received

continuous elevated N deposition rates for many

years.For one of these point sources NH3, deposition

rates ranged from 73 kg NH3-N haÿ1 yÿ1 at the wood-

land edge (30 m downwind of an intensively managed

poultry farm) to 18 kg NH3-N haÿ1 yÿ1 110 m down-

wind, and <7 kg NH3±N haÿ1 yÿ1 650 m downwind. The

elevated N deposition increased the soil mineral N

concentrations and ¯uxes of NO and N2O in a linear

fashion, and 6.3% of the NH3 deposited within the

®rst 100 m downwind of the farm was emitted as N2O

(>4 kg N2O±N haÿ1 yÿ1) (Skiba et al., 1998). High N2O

emissions were also reported for a spruce forest in

382 U. Skiba, K.A. Smith / Chemosphere ± Global Change Science 2 (2000) 379±386

Germany receiving N deposition rates of 40 kg N haÿ1

yÿ1. Mean monthly N2O emissions ranged from 4 to

16 lg N mÿ2 hÿ1, but in a comparable spruce forest in

Ireland with low deposition rates (<10 kg N haÿ1 yÿ1),

N2O emissions were only 6 2 lg N mÿ2 hÿ1 (Butter-

bach-Bahl et al., 1998). In a forest in South Germany,

where N deposition rates were <15 kg N haÿ1 yÿ1, N2O

was mainly deposited (Rennenberg et al., 1998). Equally

in phosphorus-limited tropical forest soils the 1% IPCC

emission factor may be an underestimate. Studies have

shown that these soils are particularly sensitive to in-

creased atmospheric inputs of N and may respond to

even small initial N additions with larger N2O emission

losses than predicted by models developed for temperate

forests (Hall and Matson, 1999).

In Fig. 1(b) many of the values below the IPCC de-

fault line were data from ®eld experiments (white dots),

where elevated N depositions were simulated. Here the

N added was largely immobilised, with N2O losses not

being signi®cantly di�erent from the control plots, to

which no N was added. The exception in this group of

experimental data was an experimental spruce planta-

tion grown on a previously agricultural soil, where the

fractional N2O loss was 4.2% (2.5 kg N2O±N haÿ1 yÿ1),

presumably due to the previously high N inputs by ag-

ricultural management (Skiba et al., 1998). Thus simu-

lation experiments of N deposition do not always mimic

the ÔrealÕ situation, and care needs to be taken when

results are interpreted.

As for agricultural soils inter-annual and seasonal

variations in the N2O ¯ux can be large and can con-

tribute to the deviations from the IPCC Ôdefault valueÕ.Seasonal variations in temperature and soil water con-

tent can usually explain part of this variability, with

maximum emissions measured at relatively high soil

moisture contents in several forests in Germany

(Schmidt et al., 1988; Brumme and Beese, 1992).

3.3. Land use management

Nitrous oxide emissions induced directly by mineral

and organic N inputs to agricultural soils and indirectly

via N deposition to natural soils are accounted for by

IPCC (1997). However, many other processes can in-

crease the soilÕs mineral N content and thereby stimulate

N2O emissions. These include land use change from

forestry to agriculture (Keller et al., 1993), biomass

burning (Weitz et al., 1998) autumn ploughing of arable

land and, as already discussed, soil drying/wetting and

freeze/thaw cycles (Davidson et al., 1993; Kaiser et al.,

1998).

All these processes stimulate N mineralisation and

temporarily reduce competition between plant and mi-

crobe for N, and thereby potentially increase emissions

of N2O, both in tropical and temperate climates. The

conversion of tropical forests to agricultural land and

biomass burning in tropical areas undoubtedly have the

largest impact on global N2O emissions. Between 1990

and 1995 the annual forest loss in developing countries

was 14 million ha. Studies in South and Central America

have shown that, in the short-term, conversion of

tropical forests to pasture substantially increases emis-

sions of N2O. In Costa Rica emissions from young

pasture (2±10 y) (380±580 lg N2O±N mÿ2 hÿ1) greatly

exceeded those from the primary forest (50±100 lg

N2O±N mÿ2 hÿ1) (Keller et al., 1993). In contrast, an old

cattle-grazed pasture emitted only one-®fth as much N2O

as that from the primary forest during the wet season

(8±12 lg N2O±N mÿ2 hÿ1 and 50 lg N2O±N mÿ2 hÿ1,

respectively), and during the dry season the older pas-

ture was a small sink for N2O (Verchot et al., 1999). In

young pastures, mineralisation of forest-derived organic

matter yields high concentrations of mineral N, which is

nitri®ed, denitri®ed and a proportion emitted as N2O,

whereas in older pastures this source of mineral N is

depleted and internal N cycling is much slower than in

the natural forest (Verchot et al., 1999).

There is very little information on the e�ect of land

use change on emissions of N2O from tropical regions of

Asia and Africa. Equally, the clear-felling of forest

plantations for timber production in temperate climates

has not received much attention. A study in northern

England, however, has shown that clear-felling increased

denitri®cation rates 10-fold: from 1±3 kg N haÿ1 yÿ1 in

the standing forest, to 10±40 kg N2O±N haÿ1 yÿ1 for the

2 years after clear-felling; N2O accounted for at least

50% of the denitri®cation end-products (Dutch and

Ineson, 1990).

Biomass burning is a source of N2O during the ®re,

and increases soil N2O emissions afterwards by Ôfertil-

isingÕ the soil by stimulating N mineralisation, and

temporarily removing the competition between plant

and microorganisms in favour of the latter. Many

forests are cleared by burning, either as an intended

permanent land-use change, or in shifting cultivation

systems. Burning is also used to rejuvenate old pasture;

it is estimated that 40% of all savannahs are burned

each year (Yienger and Levy, 1995). In a burned semi-

arid chaparral in California, USA, burn-induced N2O

emissions were long-lasting and persisted for at least 6

months (Levine et al., 1988). In Costa Rica a secondary

forest was cleared, which increased N2O emissions

from 15 to 27 lg N2O±N mÿ2 hÿ1, and then burned,

inducing a sharp short-lived emission peak of 1230 lg

N2O±N mÿ2 hÿ1 for 3 days, which then declined. Post-

burn N2O emissions were elevated for 3±4 months after

the event, with mean emissions of 175 lg N2O±N mÿ2 hÿ1

(Weitz et al., 1998). Surprisingly, for a savannah soil

in South Africa, N2O emissions were not detected,

even after burning and wetting events (Levine et al.,

1996).

U. Skiba, K.A. Smith / Chemosphere ± Global Change Science 2 (2000) 379±386 383

3.4. Temperature

Like any biological process, rates of nitri®cation and

denitri®cation increase with increasing the temperature.

Direct linear relationships between N2O emission and

seasonal and diurnal temperature changes have been

shown for many soils in temperate climates. For N2O

emissions from a range of acid forest soils in Scotland,

activation energies ranged between 70 and 170 kJ molÿ1

(Skiba et al., 1998). In agricultural soils in Scotland, it

has been shown that if soil WFPS or mineral N content

are limiting, there may not be a very clear relationship

with temperature. However, when only those data

points where the other factors are non-limiting are

considered, there is evidence of very steep responses to

temperature, with Q10 values of up to 8 (Dobbie et al.,

1999).

In tropical natural soils, where seasonal variations in

temperature are much smaller, evidence of diurnal

variations is mixed. For example, in the closed canopy

of a terra ®rme forest in Brazil, no diurnal variations in

N2O emission were observed (Matson et al., 1990), but,

in a semi-deciduous forest in Venezuela day-time ¯uxes

were typically 50% larger than night-time ¯uxes. How-

ever, in a nearby savannah diurnal temperature changes

did not a�ect N2O emissions (Sanhueza et al., 1990).

3.5. Soil water content

For many soils soil water content (through its e�ect

on aeration), together with N supply, has been shown to

be the dominant variable controlling the N2O emission

rate; to some extent this has already been discussed in

the previous sections. In order to improve current

methodologies for estimating N2O emissions, inter-

annual variations in rainfall intensity, frequency and

timing, particularly in relation to fertiliser applications,

need to be taken into account. Davidson (1991) has

produced a model of the relationship between the water-

®lled pore space (WFPS) of the soil and N trace gas

emissions, which suggests N2O emissions are at a max-

imum at a WFPS of 60%, with nitri®cation and deni-

tri®cation contributing about equally towards its

production. However, later studies, both in the tropics

(Veldkamp et al., 1998) and in a temperate climate

(Dobbie et al., 1999) suggest that maximum N2O emis-

sions occur at a WFPS of 80±85% (Fig. 2). Further

support for the importance of this higher range comes

from the work of Ruser et al. (1998) which indicated

that the highest ¯uxes were induced by the loss of

macro-pores due to compaction, which increased WFPS

to a mean value of 85%. These data suggests that much

wetter and a much greater degree of anaerobicity is re-

quired to produce maximum N2O emissions than pre-

viously suggested by (Davidson, 1991).

4. Summary and outlook for future work

Nitrous oxide emission factors vary widely from

0.1% to 7% in both agricultural and natural soils. Rea-

son for this deviation were inter-annual variations due

to variations in rainfall, timing and intensity and the

contribution of N2O by indirect sources i.e. ploughing,

winter-time emissions, excessive emissions from forest

soils in high N deposition areas. The importance of

rainfall and water ®lled pore space on variations of N2O

emissions suggests that the uncertainty in budgets can

only be alleviated by including these variables into the

budget equation, ideally in multilayered models rather

than simple emission factor style.

Acknowledgements

The authors wish to thank the organisers and spon-

sers of the International Workshop on Atmospheric

N2O emissions in Tsukuba, Japan, March 1999 for in-

viting us to contribute and for providing ®nancial sup-

port for Ute Skiba to attend the meeting.

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Dr. Ute Skiba is a Senior Scienti®c O�cer at ITE, specialising inmeasurements and scaling of ¯uxes of N2O and NO from ag-ricultural and semi-natural soils.

Prof. Keith Smith is a Professorial Fellow in IERM. His work isfocussed on trace gas exchange, in particular the measurementand modelling of N2O emissions from agricultural soils,methane emission from rice ®elds and oxidation of atmosphericmethane in soils.

386 U. Skiba, K.A. Smith / Chemosphere ± Global Change Science 2 (2000) 379±386